Ecological Risk Assessment for Contaminated Sites - Chapter 4 potx

4 Analysis of Effects What is there that is not poison? All things are poison, and nothing is without poison. Solely the dose determines that a thing is not a poison. —Paracelsus, translation by Deichmann et al. (1986) In the analysis of effects, assessors determine the nature of toxic effects of the contaminants and their magnitude as a function of exposure. Effects data might be available from field monitoring, from toxicity testing of the contaminated media, and from traditional single-chemical laboratory toxicity tests (Table 4.1). The asses- sor must evaluate and summarize the data concerning effects in such a way that it can be related to the exposure estimates, thereby allowing characterization of the risks to each assessment endpoint during the risk characterization phase. In the analysis of effects, available effects data must be evaluated to determine which are relevant to each assessment endpoint, and they must be reanalyzed and summarized as appropriate to make them useful for risk characterization. Two issues must be considered. First, what form of each available measure of effect best approximates the assessment endpoint? This issue should have been considered during the problem formulation. However, the availability of unanticipated data and better understanding of the situation after data collection often require reconsider- ation of this issue. The second issue in analysis of effects is expression of the effects data in a form that is consistent with expressions of exposure. Integration of exposure and effects defines the nature and magnitude of effects, given the spatial and temporal pattern of exposure levels. Therefore, the relevant spatial and temporal dimensions of effects must be defined and used in the expression of effects. For example, if the exposure is to a material such as unleaded gasoline that persists at toxic levels only briefly in soil, then effects that are induced in that time period must be extracted from the effects data for the chemicals of concern, and the analysis of field-derived data should focus on biological responses such as mass mortalities that could occur rapidly rather than long-term average properties. The degree of detail and conservatism in the analysis of effects depends on the tier of the assessment (Chapter 1). Scoping assessments need only determine qual- itatively that an effect may occur because a receptor is potentially exposed to one or more contaminants. Screening assessments typically define the exposure–effects relationship in terms of a benchmark value, a concentration that is conservatively defined to be a threshold for toxic effects (Chapter 5). Definitive assessments should define the exposure–response relationship (Chapter 6) and should separately estimate the uncertainty concerning that relationship (Chapter 7). © 2000 by CRC Press LLC 4.1 SINGLE-CHEMICAL OR SINGLE-MATERIAL TOXICITY TESTS In ecological risk assessments for contaminated sites, single-chemical or single- material (e.g., gasoline) toxicity data are usually obtained from the literature or from databases rather than generated ad hoc. One source is the EPA ECOTOX database, which contains toxicity data for aquatic biota, wildlife, and terrestrial plants. It is available from the EPA and commercial sources (http://hepa.gov/medectox). Assessors must select data that are most relevant to the assessment endpoints and that can be used with the exposure estimates. As far as possible, data should be selected to correspond to the assessment endpoint in terms of taxonomy, life stage, response, exposure duration, and exposure conditions. However, because the vari- ance among chemicals is greater than the variance among species and life stages, any toxicity information concerning the chemicals of interest is potentially useful. If no toxicity data are available that can be applied to the assessment endpoints (e.g., no data for fish or no reproductive effects data), or if the test results are not applicable to the site because of differences in media characteristics (e.g., pH or water hardness), tests may be conducted ad hoc. However, most tests performed for specific sites are tests of local contaminated media (Section 4.2) rather than single chemicals. If combined toxic effects of multiple contaminants are thought to be significant, and if appropriate mixtures are not available in currently contaminated media, synthetic mixtures may be created and tested. Toxicity tests of single chemicals that are obtained from published literature have biases that should be understood by ecological risk assessors. Assessors must be aware of these biases when these data are used to derive toxicity benchmarks or exposure–response models for chemicals. Potential sources of bias in the test data include the following: • The forms of chemicals used in toxicity tests are likely to be more toxic than the dominant forms at hazardous waste sites. For metals the tested forms are usually soluble salts, and organic chemicals may be kept in aqueous solution by cosolvents. In dietary or oral dosing tests organic chemicals are typically dissolved in readily digested oils. TABLE 4.1 Types of Effects Data Used in Ecological Risk Assessments of Contaminated Sites and Sources of the Data Type Source Single-chemical toxicity Published scientific literature reporting results of toxicity tests with individual chemicals or materials and summarizations of that literature such as water quality criteria Ambient media toxicity Site-specific in situ or laboratory toxicity tests of contaminated water, sediment, soil, or food Biological survey Site-specific sampling or observation of organisms, populations, or communities in contaminated areas © 2000 by CRC Press LLC • Combined toxic effects are not observed in toxicity tests of single chemicals. • The test species for toxicity testing may not be representative of the sensitivity of species native to the site. • The standard media used in toxicity tests may not be representative of those at a particular contaminated site. For example, aqueous tests typi- cally use water with moderate pH and hardness with little suspended or dissolved matter, and soil tests typically use agricultural loam soils or artificial soils. • Laboratory test conditions may not be representative of field conditions (e.g., temperature, use of sieved soil, and maintenance of constant moisture). 4.1.1 T YPES OF T OXICITY T ESTS Conventionally, toxicity tests determine effects on individual organisms and are divided into two classes, acute and chronic. Acute tests are those that last a small proportion of the life span of the organism (<10%) and involve a severe effect (usually death) on a large proportion of exposed organisms (conventionally, 50%). Acute tests also usually involve well-developed organisms rather than eggs, larvae, or other early life stages. Chronic tests include much or all of the life cycle of the test species and include effects more subtle than death (e.g., reduced growth and fecundity). In these tests, the endpoint is typically based on statistical significance, so the proportion affected may be large or small. In addition, there are many tests that fall between these two types, which are termed subchronic, short-term chronic, etc. They typically have short durations but include sublethal responses. A prominent example is the 7-day fathead minnow test, which includes growth as well as death. This test includes only part of one life stage but uses the larval rather than juvenile stage and statistical significance rather than effects levels to derive the test endpoint (Norberg and Mount, 1985). In general, tests with longer durations, more life stages, and more responses reported are more useful for risk assessment, because they provide more information and because the exposures at contaminated sites are typically chronic. However, if exposures are acute, then acute tests should be preferred. Examples include expo- sures of transients such as migratory waterfowl or highly mobile species that may use a site in transit or exposures during episodes of contamination, such as overflow of waste ponds or flushing of contaminants into surface waters by storms. Following are general recommendations for selecting toxicity tests of single chem- icals or materials. Other issues specific to tests of particular media are addressed later. Standardization — In general, choose standard tests. Standard test protocols have been developed or recommended by governments (Keddy et al., 1995; EPA, 1996a) and standards organizations (OECD, 1998; APHA, 1999; ASTM, 1999). Most extrapolation models for relating test endpoints to assessment endpoints require standard data (Section 4.1.9.1). In addition, results of standard tests are likely to be reliable because of the QA/QC procedures that are part of the standard methods, and because test laboratories are likely to conduct standard tests routinely. However, nonstandard tests should be used when particular site-specific issues cannot be resolved by standard test results. © 2000 by CRC Press LLC Duration — Choose tests with appropriate durations. Two factors are relevant. The first is the duration of the exposures in the field. If exposures are episodic, as is often the case for aqueous contamination, then tests should be chosen with durations as great as the longest episodes. The second factor is the kinetics of the chemical. Some chemicals such as chlorine in water or low-molecular-weight nar- cotics are taken up and cause death or immobilization in a matter of minutes or hours. Others such as dioxins have very slow kinetics and require months or years to cause some effects such as reproductive decrements. In general, longer durations (i.e., chronic tests) are preferred, but these site-specific considerations may override that generality. Response — Choose tests with appropriate responses. In particular, if an appar- ent effect of the contaminants has been observed in field studies, tests that include that effect as a measured response should be used. More generally, chosen tests should include responses that are required to estimate the assessment endpoint. Since most tests are of collections of organisms, and assessment endpoints are usually defined at the population or community level, choose responses that are relevant to higher levels of organization including mortality, fecundity, and growth. Physiolog- ical and histological responses are generally not useful for estimating risks, because they cannot be related to effects at higher levels. However, if they are characteristic of particular contaminants, they may be useful for diagnosis (Chapter 6). Consistency — Prefer tests matching ambient media tests performed at the site (Section 4.2). For example, if 7-day fathead minnow larval tests are performed with ambient water, use of the same tests with individual chemicals can help in interpre- tation of results. Media — Prefer tests conducted in media with physical and chemical properties similar to the site media. Organisms — Prefer taxa and life stages that are closely related taxonomically to the endpoint species. If an assessment endpoint is defined in terms of a community, one may either choose tests of species that are closely related to members of the community or use all high quality tests in the hope of representing the distribution of sensitivity in the endpoint community (Section 4.1.9.1). Species, life stages, and responses should also be chosen so that the rate of response is appropriate to the duration of exposure and kinetics of the chemical. In general, responses of small organisms such as zooplankters and larval fish are more rapid because they achieve a toxic body burden more rapidly than larger organisms. Therefore, if exposures are brief and if those small organisms are relevant to the assessment endpoint, tests of small organisms should be preferred over larger organisms that are no more relevant. However, such tests may not be appropriate if, for example, the endpoint is fish kills, or exposures do not occur during the breeding period of fish. Multiple exposure levels — Studies that employ only a single concentration or dose level plus a control are seldom useful. If the exposure causes no effect, it may be considered a no observed effects level (NOEL), but no information is obtained about levels at which effects occur. Conversely, if the exposure causes a significant effect, it may be considered a lowest observed effects level (LOEL), but the threshold for effects cannot be determined. Studies in which multiple exposure levels were applied allow an exposure–response relationship to be evaluated and NOELs and © 2000 by CRC Press LLC LOELs to be determined. Consequently, studies that apply multiple exposure levels are strongly preferred. Exposure quantification — To interpret the results of toxicity tests correctly and to apply these results in risk assessments, the exposure concentrations or doses should be clearly quantified. Ideally, the test chemical should be measured at each exposure level; measured concentrations are always preferable to nominal concentrations. Chemical form — Correct estimation of the dose requires that the form of toxicant used in the test be clearly described. For example, in tests of lead, the description of the dosing protocol should specify whether the dose is expressed in terms of the element (e.g., lead) or the applied compound (e.g., lead acetate). Tests of chemicals in the forms occurring on the site should be preferred. This is partic- ularly important for chemicals that may occur in multiple forms under ambient conditions that have widely differing toxicities. Statistical expressions of results — The traditional toxicity test endpoints for chronic tests, NOELs and LOELs, have low utility for risk assessment (Suter, 1996a). NOELs are the highest exposure levels at which no effects are observed to differ statistically significantly from controls, while LOELs are the lowest exposure levels at which one or more effects are observed to differ statistically significantly from controls. These endpoints do not indicate whether the statistically significant effect is, for example, a large increase in mortality or a small decrease in growth. The level of effect at a NOEL or LOEL is an artifact of the replication and dosing regime employed. Use of the NOEL or LOEL does not indicate how effects increase with increasing exposure, so the effects of slightly exceeding a NOEL or LOEL are not qualitatively or quantitatively distinguishable from those of greatly exceeding it. To estimate risks, it is necessary to estimate the nature and magnitude of effects that are occurring or could occur at the estimated exposure levels. To do this, exposure–response relationships should be developed for chemicals evaluated in ecological risk assessments. Methods for fitting of exposure–response distributions to toxicity data are presented in Crump (1984), Kerr and Meador (1996), Moore and Caux (1997), and Bailer and Oris (1997). In some cases these criteria may conflict. Hence, assessors must determine their relative importance to the particular site and assessment, and apply them accordingly. 4.1.2 A QUATIC T ESTS More toxicity tests are available for aquatic biota than any other type of receptor. In general, flow-through tests are preferred over static-renewal tests, which are preferred over static tests. Flow-through tests maintain constant concentrations, whereas concentrations may decline significantly in static tests. However, in a few cases, static tests are appropriate, because exposure is static, as in the spillage of a chemical into a pond. The most abundant type of test endpoint is the 48- or 96-h LC50. However, chronic test results are more generally useful. They include life cycle tests and, for fish, early life-stage tests. Currently, the most popular aquatic test organisms in the United States are fathead minnows ( Pimephales promelas ) and daphnids ( Daphnia and Ceriodaphnia spp.). Test results for algae or other aquatic plants are less often available. Aquatic microcosm and mesocosm test data are rare, and largely limited to pesticides. © 2000 by CRC Press LLC 4.1.3 S EDIMENT T ESTS Selecting representative sediment tests and test results is complicated by the inter- actions among the multiple phases (i.e., particles, pore water, and overlying water) of the sediment system. Chemicals can be tested either in the presence (spiked- sediment tests) or absence (aqueous) of the solid phase. Test selection depends on the expected mode(s) of exposure, and more than one test type may be appropriate. Spiked-sediment tests consist of the addition of known quantities of the test chemical or material to a natural or synthetic sediment to which the test organism is exposed. Spiked-sediment tests provide an estimate of effects based on all direct modes of exposure, including ingestion, respiration, and absorption. Hence, toxicity to sedi- ment-ingesting organisms may be best approximated by bulk sediment tests. The primary disadvantage is that the exposure–response relationship is somewhat uncer- tain due to the unquantified effects of the sediment matrix (Ginn and Pastorok, 1992). Aqueous phase tests are most appropriate if interstitial or overlying water is believed to be the primary exposure pathway for the toxicants and receptors at a site. As noted in Section 4.1.2, aqueous tests and data are more abundant than any other kind. Most of the species tested live in the water column rather than the sediment. Aqueous tests and data are used to evaluate aqueous exposures of benthic species, based on data suggesting that benthic species are not systematically more sensitive than water column species (EPA, 1993a). The types of aqueous tests and factors to consider in selecting a test type are discussed in Section 4.1.2 and apply here as well. Sediment and water tests are available for marine and freshwater species (Section 4.2.2). Risk assessors should choose tests in media similar to the site media. Unlike aqueous toxicity data, which are relatively abundant for both fresh water and salt water, there are few test data from freshwater sediment tests relative to estuarine sediment tests. Therefore, it is necessary to consider whether to use saltwater toxicity values for assessments of freshwater systems. Klapow and Lewis (1979) applied a statistical test of medians to freshwater and marine acute toxicity data for nine heavy metals and nonchlorinated phenolic compounds. In only one case (cadmium) was there a statistically significant difference in the median response of marine and freshwater organisms. On the other hand, Hutchinson et al. (1998) found potentially important differences. They compared the aqueous toxicity of several heavy metals, pesticides, and organic solvents to freshwater and saltwater invertebrates—83% of the no observed effects concentrations (NOEC) and 33% of the 50% effects con- centrations (EC50) for freshwater and saltwater invertebrates were within a factor of 10. Based on the ratios of EC50s, freshwater invertebrates were more sensitive than saltwater invertebrates to four (2-methylnaphthalene, 1-methylnaphthalene, benzene, and chromium) of the 12 evaluated chemicals. Comparison of NOECs indicated that two (copper and cadmium) of the six chemicals for which sufficient data were available to allow comparison were more toxic to freshwater invertebrates than to saltwater invertebrates. The authors emphasized that the results should be considered preliminary because of the limited amount of appropriate data. The bottom line is that cautiously using data from tests of saltwater sediments to evaluate chemicals in freshwater sediments is probably better than having no data at all in © 2000 by CRC Press LLC the preliminary stages of an assessment. There is precedent for this in the use of effects range–low values from estuarine and marine sediments (Long et al., 1995) as ecotox thresholds for both marine and freshwater sediments (Office of Emergency and Remedial Response, 1996). The physical and chemical properties of the test media are particularly important for evaluating chemical toxicity in the sediment system. Characteristics of the sedi- ment (e.g., organic carbon content and grain size distribution) and water (e.g., dis- solved organic carbon, hardness, and pH) can significantly alter the speciation and bioavailability of the tested material. Again, tests in media similar to the site media should be preferred. Regression models could be derived to account for confounding matrix factors (e.g., grain size or organic carbon content) (Lamberson et al., 1992). However, such models are species- and matrix factor-specific and would need to be developed on a case-by-case basis. This is not practical for most hazardous waste site assessments, especially for adjustments of multiple variables. The test method also can affect exposure. For example, chemical concentrations and bioavailability can be altered by the overlying water turnover rate, the water-to-sediment ratio, and the oxygenation of the overlying water (Ginn and Pastorok, 1992). Issues associated with sediment toxicity testing are discussed in detail elsewhere (Burton, 1992). 4.1.4 S OIL T ESTS The available body of soil toxicity tests is relatively small and poorly standardized. For example, few organic chemicals other than pesticides are represented. Soil toxicity test data for inorganic chemicals and some organic compounds are available for plants (mainly crops), soil invertebrates (primarily earthworms), and soil micro- organisms (usually expressed as changes in rates of carbon mineralization, nitrifi- cation, nitrogen fixation, or other processes). Tests in both soil and soil solution may be useful for assessing risks from soil contaminants. The relevance of published tests in soil to the assessment of risks to soil organisms seems self-evident, but, unless the properties of the test soil are similar to those of the site soil, the toxicity observed in the test soil concentration may be poorly correlated with effects at the site. For example, Zelles et al. (1986) found effects of chemicals on microbial processes to be highly dependent on soil type. Moreover, it is usually desirable for the assessor to exclude data from tests in quartz sand or vermiculite, unless toxicity of chemicals mixed with these mate- rials is demonstrated to be similar to that in natural soils. Tests conducted in solution have potentially more consistent results than those conducted in soil. Toxicity observed in inorganic salts solution may be related to concentrations in soil extracts, estimated pore water concentrations, or springs where wetland plant communities are located. It has even been proposed that aquatic toxicity test results could be used to estimate the effects of exposure of plants and animals to contaminants in soil solution (van de Meent and Toet, 1992; Lokke, 1994), although we do not recommend this practice. The risk assessor should be aware that bioavailability in soil from the contam- inated site may be substantially different from the bioavailability in published soil tests. As stated in Section 3.4.1, aged organic chemicals are typically less available © 2000 by CRC Press LLC and less toxic to biota than organic chemicals freshly added to soil in published toxicity tests (Alexander, 1995); thus, the toxicity at the contaminated site may be overestimated if a published toxicity test of a chemical freshly added to soil is emphasized too heavily in the assessment. The risk assessor can make adjustments to observed toxic concentrations to account for differences in soils or chemical speciation. The variance in toxicity among natural soils may be reduced by normal- izing the test soil concentrations to match normalized site soil concentrations (Sec- tion 3.4.1.1). Or free metal activities in soil solution may be estimated, potentially improving the precision of toxic thresholds for plants, soil invertebrates, or microbial processes (Sauvé et al., 1998). The assessor may be more liberal in including tests in screening assessments (e.g., in the derivation of screening benchmarks) than in definitive assessments. In definitive assessments, soil type and chemical speciation should be factors in decisions about the acceptability of data. Tests should be chosen for risk assessments based on a relationship to the assessment endpoint. For example, if the assessment endpoint is production of the plant community, tests relating to plant growth or yield or mycorrhizal biomass may be sufficiently relevant to the endpoint, but tests of DNA damage would probably not be. Tests of litter-feeding earthworms may not be representative of those that ingest soil, and vice versa. Similarly, it is not always clear that microbial communities that have become altered in their tolerance of contaminants (pollution-induced com- munity tolerance, PICT; Rutgers et al., 1998) are indicators of a decrease in the rate of a valued microbial process (Efroymson and Suter, 1999). Microcosm tests of the soil community and processes such as decomposition incorporate indirect effects of chemical addition as well as direct toxic effects (Sheppard and Evenden, 1994; Bogomolov et al., 1996; Parmelee et al., 1997; Salminen and Sulkava, 1997; Weeks, 1998). In addition, in microcosms, the responses of communities may be observed directly rather than deduced from effects on single populations of invertebrates. The assessment endpoint may include a defined level of effects such as a 20% reduction in some endpoint property. However, such a decrease in the rates of some microbial processes such as litter decomposition may be desirable (or acceptable) in particular ecosystems (Efroymson and Suter, 1999); thus, an appropriate level of effects is sometimes unclear. Moreover, the desired level of effect is seldom obtain- able from soil toxicity test results in the literature. Frequently, the EC50 is reported, but lower-level or lower-percentile effects are not. Often the lowest observed adverse effects level is a 50% effects level or higher, and lower concentrations (other than the reference) were not tested. No good models for estimating an EC20 from an EC50 exist for plants, earthworms, or other soil organisms. For example, the shape of the dose–response curve may be affected by whether the chemical is an essential element or whether detoxification occurs. It is advisable for the assessor either to use a safety factor or retain the uncertainty associated with the single-chemical toxicity test line of evidence during the risk characterization (Section 4.1.9.2). 4.1.5 D IETARY AND O RAL T ESTS Dietary and oral toxicity tests are those in which test animals are exposed to toxicants orally in food, water, or another carrier, with the organ of uptake being the © 2000 by CRC Press LLC gastrointestinal tract. These tests are employed primarily with birds and mammals and are rarely applied to aquatic organisms. For dietary tests, the toxicant is mixed with food or water and test animals are allowed to feed ad libitum . The amount of food consumed daily should be recorded so that the daily dose can be estimated. A potential problem with dietary tests is that animals may not experience consistent exposure throughout the course of the study. For example, as animals become sick, they are likely to consume less food and water. They may also eat less or refuse to eat if the toxicant imparts an unpleasant taste to the food or water or if the toxic effects induce aversion. In oral tests, animals receive periodic (usually daily) toxicant doses by gavage (i.e., esophageal or stomach tube) or by capsules. The chemical is generally mixed with a carrier (e.g., water, mineral oil, acetone solution, etc.) to facilitate dosing. Oral tests assure consistent daily doses of test chemicals including those that are repellent or aversive. The choice of carrier used for oral or dietary tests has been shown to influence uptake by binding with the toxicant or otherwise influencing its absorption. For example, Stavric and Klassen (1994) report that the uptake of benzo( a )pyrene by rats is reduced by both food and water but facilitated by oil. Similarly, uptake of inorganic chemicals varies dramatically between tests with food and with water as carriers. Chemicals are generally taken up more readily from water than from food. Results of most dietary toxicity tests are presented as toxicant concentrations (mg/kg) in food or water. These data can then be converted into doses (mg toxicant/kg body weight/day) by multiplying the concentrations in food or water by food ingestion rates and body weights either reported in the literature or presented in the study (e.g., Sample et al., 1996a). Fairbrother and Kapustka (1996) argue that uncertainty in food consumption rates, particularly in response to toxicity, precludes the accurate esti- mation of dose, and therefore concentration data should not be converted to dose. They are correct in indicating that the conversion is a significant source of uncertainty. However, toxicity data expressed as concentrations cannot be readily compared to multimedia contaminant exposure estimates (Section 3.10). Therefore, the conversion of concentration to dose is recommended, unless only one source of exposure is significant. Conversion of results from most oral toxicity tests is not needed as the results are generally expressed as dose in mg/kg/day or equivalent metrics. Standard methods for performing avian and mammalian oral toxicity tests have been developed and are generally applied for testing of drugs, pesticides, and other chemicals. While standard test methods specifically developed for wildlife at haz- ardous waste sites do not exist, existing standard laboratory tests may be modified and applied. These tests vary from acute tests to subacute dietary tests to develop- mental and reproductive tests. A summary of selected standard oral test methods is presented in Table 4.2. 4.1.6 B ODY B URDEN –E FFECT R ELATIONSHIPS Single-chemical toxicity tests may be used to develop exposure–response relation- ships based on internal exposure measures (body burdens), rather than on external exposures (media concentrations or administered doses). In theory, this approach © 2000 by CRC Press LLC TABLE 4.2 Selected Standard Oral Toxicity Methods for Birds and Mammals Taxon Test Type Test Species Duration Exposure Route(s) Test Endpoint(s) Ref. Mammal Acute Not stated Single dose, 14 day post Gavage Mortality ASTM, 1999 Rats Single dose, 7 day post Gavage Mortality ASTM, 1999 Subacute dietary Rats 90 day Capsule, gavage, in diet, in water Mortality, organ pathology, behavior ASTM, 1999 Developmental Rats, rabbits Day 6–15 of gestation (rats) Day 6–18 of gestation (rabbits) Capsule, gavage Fertility, fetal body weights, number of dead fetuses, number of malformed fetuses ASTM, 1999 Bird Subacute dietary Northern bobwhite, Japanese quail, mallard, ring-necked pheasant 5 day exposure, 3 day post In diet Mortality, but other effects can also be considered ASTM, 1999 Reproduction Northern bobwhite, mallard 10 week In diet Adult mortality, eggs laid, egg fertility, egg hatchability, eggshell thickness, weight and survival of young ASTM, 1999, EPA, 1991b © 2000 by CRC Press LLC © 2000 by CRC Press LLC [...]... Variable Perciformes Tetraodontiformes Perciformes Gasterosteiformes Atheriniformes Cypriniformes Tetraodontiformes Perciformes Gasterosteiformes Tetraodontiformes Perciformes Tetraodontiformes n 190 12 34 8 46 7 46 148 36 8 33 34 95% PI 63 13 25 16 9 501b 13 25 20 20 32 25 23.5 26.0 99% PI 83 18 34 24 12 786b 17 33 27 30 43 34 31.7 34. 5 Uncertainty factors calculated by Calabrese and Baldwin (19 94) ; used... 15 43 TABLE 4. 7 (continued) Taxonomic Extrapolation: Means and Weighted Means Calculated for the 95% and 99% Prediction Intervals (PIs) for Uncertainty Factors Calculated from Hierarchical Regressionsa Uncertainty Factor X Variable Siluriformes Anguiliformes Anguiliformes Anguiliformes Anguiliformes Atheriniformes Atheriniformes Atheriniformes Atheriniformes Gasterosteiformes Gasterosteiformes Perciformes... 28 89 42 26 37 14 90% 26 18 18 32 26 28 91 34. 1 27.1 95% 50 32 31 64 50 53 246 75.1 54. 7 99% 198 106 93 228 206 188 2 247 46 6.6 2 64. 9 Regression analysis from Suter et al (1987) Decrease in weight of fish at end of larval stage Source: Calabrese, E J and Baldwin, L A., Performing Ecological Risk Assessments, Lewis Publishers, Boca Raton, FL, 1993 With permission b © 2000 by CRC Press LLC TABLE 4. 7 Taxonomic... Cichlidae 5 13 Esocidae 11 9 Cyprinodontidae 32 7 Labridae 12 55 Poecillidae 12 3 14. 4 12.6 14 6 24 13 9 78 5 21.3 17.9 Salmoniformes Salmoniformes Salmoniformes Cypriniformes Cypriniformes Taxonomic Extrapolation: Orders within Classes Cypriniformes 225 Siluriformes 203 Perciformes 44 3 Siluriformes 111 Perciformes 219 Oncorynchus Oncorynchus Salmo Carassius Carassius Cyprinus Lepomis Lepomis Cyprinodon... widely used for the development of toxicity values for human health risk assessment (Dourson and Stara, 1983) and have also been applied for wildlife risk assessment (e.g., Banton et al., 1996; Sample et al., 1996b; Hoff and Henningson, 1998) Uncertainty factors have been applied to account for a wide range of extrapolations (e.g., interspecies, acute-to-chronic, laboratory-to-field, LOAEL-to-NOAEL) While... routinely engage in ecological risk assessments of contaminated sites often develop standard ecotoxicity profiles (Sample et al., 1997b) These off-the-shelf profiles are labor saving and also are useful for planning the assessment during the problem formulation process (Chapter 2) Reviews such as those produced by R Eisler for the National Biological Service may also be useful during problem formulation (www.pwrc.usgs.gov/new/chrback.htm)... benchmarks He found that it was a reasonable threshold value for PAH effects in independent data sets from PAH -contaminated sites 4. 1.9 SINGLE-CHEMICAL TEST ENDPOINTS AND DEFINITIVE ASSESSMENT Single-chemical toxicity test endpoints can play two roles in definitive assessments If biological surveys or ambient media toxicity tests are performed, single-chemical test results may be used to support the conclusion... adequate for screening purposes, but not for definitive risk characterization For definitive assessments, such nonspecific distributions can be parsed into distributions of thresholds for specific effects For example, Jones et al (1999) developed distributions of community-level effects and lethality from the sediment toxicity data presented in McDonald et al (19 94) and Long and Morgan (1991) As in Figure 4. 1,... TABLE 4. 5 Linear Equations for Extrapolating from Standard Fish Test Species to All Freshwater or Marine Fish (units are log µg/l) Test Species Pimephales promelas Lepomis macrochirus Oncorhynchus mykiss Cyprinodon variegatus Slope 1.01 Intercept –0.30 n 3 54 mean X 2.77 F1 0 .45 F2 0.0006 PIa 1.31 0.96 0.17 500 2.52 0 .49 0.0005 1.37 0.99 0.29 48 0 2 .42 0.38 0.00 04 1.20 0.97 0.03 51 1.25 0.58 0.0085 1 .49 ... chemical of interest or for a similar chemical, they recommend generic factors of 1.2 for birds and 0. 94 for mammals Factors must be developed for chronic exposures and for effects other than lethality PBPK models — A final approach that may be used for wildlife extrapolation is PBPK models These are compartmental models that combine contaminant-specific characteristics with species-specific anatomically . 7). © 2000 by CRC Press LLC 4. 1 SINGLE-CHEMICAL OR SINGLE-MATERIAL TOXICITY TESTS In ecological risk assessments for contaminated sites, single-chemical or single- material (e.g., gasoline). readily digested oils. TABLE 4. 1 Types of Effects Data Used in Ecological Risk Assessments of Contaminated Sites and Sources of the Data Type Source Single-chemical toxicity Published scientific. useful for diagnosis (Chapter 6). Consistency — Prefer tests matching ambient media tests performed at the site (Section 4. 2). For example, if 7-day fathead minnow larval tests are performed